Abstract

The Fungicide Vinclozolin (VZ) is an endocrine disruptor with anti-androgenic
activity whose potential for disrupting epigenetic mechanisms in mammals has
also been shown. In this regard, this study combined reproductive endpoints
and effect assessments at the sub-individual level associated with Genotoxicity
and altered DNA methylation patterns. Physaacuta was exposed to VZ at
nominal concentrations from 0.0005 to 5mg/L. Fecundity (production of F1 eggs)
over 45 days and fertility (viability of F1 eggs) after further 21-dayembryonic
development under exposure and non-exposure conditions were monitored. The
Genotoxicity of VZ was evaluated by micronuclei scoring. Potential epigenetic
alterations were measured by scoring DNA methylation patterns. Although
acute exposure (96h prescreening test) at 5mg/L did not caused mortality,
long-term exposure (45-day reproduction test) affected the survival of snails.
Reduced effects on reproduction, micronuclei induction and DNA demethylation
events were observed at 5mg/L. At the other VZ concentrations no effects on
these endpoints were found. Thus reproduction endpoints were as sensitive
as the effects at the sub individual level. This reproduction test, with additional
assessments of the hatching success of the F1 recovered eggs, assessed
reliable reproduction success in freshwater snails. The use of exposed and nonexposed
F1 embryos allowed us to observe if damage due to parental exposure
could be maintained or recovered when suspending treatment during embryonic
development. Finally, the combination of long-term ecotoxicological effects and
genotoxic/epigenetic biomarkers can broaden our understanding of pollutant
impacts.

Introduction

Continuously increasing chemical pollution has led to a complex
environmental scenario in which chronic tests on a wide range of
species and specific endpoints are crucial tools for studying chemical
effects. Chronic testing can provide information on processes such
as growth, development and reproduction, which are relevant at
the population level. Aquatic invertebrates are target organisms
for different OECD chronic toxicity tests [1-4] including one lifecycle
test [5]. Although no mollusk-based toxicity tests have been
internationally standardized, an OECD project is developing a
partial life-cycle test with gastropods [6,7]. The OECD/EDTA
(Endocrine Disrupters Testing and Assessment) Conceptual
Framework has indicated the need for mollusc-based tests, and
the phylum offers species of ecological and economic relevance
that are known to be uniquely sensitive to a number of Endocrine
Disrupter Compounds (EDCs) [8,9] Some assays performed with
freshwater gastropods have addressed the effects of xenobiotics
their reproductive capacities (fecundity and fertility) after long-term
exposure. Oehlmann et al. [10] reported malformations of the female genital system, and massive stimulation of oocyte and spawning
mass production in Marisa cornuarietis induced by Bisphenol A
and Octyl phenol with a complete life-cycle. These endpoints were
also assessed on Lymnaeaacuminata after exposure to pyrethroid
pesticides [11] Czech et al. [12] exposed adults of Lymnaeastagnalis
to Tributyltin, β-sitosterol and 4-nonylphenol with a view to looking
at the reproductive and histopathological effects on the F0 and F1
generations. Leung et al. [13,14] also exposed Lymnaeastagnalis and
Physafontinalis from embryos to sexual maturity to Tributyltin.
Recently, different approaches have been developed to perform
embryo tests with freshwater snails to form part of mollusc lifecycle
tests [8, 15-18]. All these studies have explored the life-stage
specific effects of molluscs to xenobiotics, while other studies have
investigated their genotoxic responses [19].

Presence of genotoxic compounds in the environment is a
matter of concern given the transfer of effects across generations.
The detection of chemical Genotoxicity has focused mainly on a
direct change at the DNA level, such as point mutations. Cytogenetic
alterations are widely accepted for detecting potential carcinogens.

The cytogenetic assay of Micronucleus (MN) induction has been
applied successfully in ecotoxicological studies [20,21]. Exposure to
chemicals can result in genetic alterations but many pollutants do
not have the capacity to induce direct DNA damage. Non-genotoxic
carcinogens are often assumed to possess an epigenetic mode of
action which can induce heritable changes that cannot be explained
by changes in DNA sequence [22,23]. Therefore, additional molecular
mechanisms need to be considered, such as epigenetics, which has
recently become a very promising target in molecular biology.
DNA methylation has been the most extensively studied epigenetic
mechanism. The evaluation of the global methylation status and the
assessment of methylation in GC-rich regions of the genome have
been proposed for both initial toxicity assessments of a compound’s
toxicity potential and an earlier indication of its possible mechanism
of action [24]. Genetic and epigenetic mechanisms are crucial for
genome stability [25] and, regardless of the chemical-induced changes
in DNA or in gene expression; both can have knock-on effects at
higher biological organization levels. Therefore studying genotoxic/
epigenetics in Ecotoxicology can help clarify how these mechanisms
can link to responses of ecological relevance.

Vinclozolin (VZ) is a nonsystemic fungicide used to control fungal
pathogens. Concern about toxicity in mammals is due to its antiandrogenic
activity [26,27]. VZ exerts its effects most dramatically
during the development stages of animals, and ultimately result in
reproductive effects [28]. VZ has been found to produce infertility of
F1 males in multigenerational studies with rats [29]. Transgenerational
effects in mammals through epigenetic mechanisms after VZ
exposure have been observed [30]. Nevertheless, other studies have
failed to find this mechanism of action [31]. The likelihood of VZ
being released into surface waters, together with its known endocrine
disruption effect, justifies undertaking chronic exposure studies with
aquatic organisms. Martinovic et al. [32] studied the reproductive
toxicity of VZ in fish and showed a concentration-dependent
reduction in fecundity (production of fewer eggs) whereas hatching
success of the deposited eggs (fertility) was not affected. VZ did not
elicit an effect on the reproduction of the freshwater invertebrate
Daphnia magna, whereas overall DNA methylation rate consistently
decreased [33,34] previously observed an earlier sexual repose and
morphological alterations of sex organs in the males of two species of
prosobranch snails. Nevertheless, it was not possible to compare these
data with fecundity data since spawning did not occur. Effects of VZ
on the reproduction of the freshwater snail Lymnaeastagnalis were
subsequently investigated by Ducrot et al., [35] who showed reduced
in fecundity (cumulated number of eggs produced per individual)
with a significant number of non-fertilized eggs as observed
microscopically. Viability of eggs was not determined. Theseauthors
recommended the complementary assessment of hatching success
in the offspring of exposed snails (fertility) in the case of fecundity
impairment when studying endocrine disruptors which may have
epigenetic effects.

The present study evaluated the fecundity and fertility of the
freshwater snail Physaacuta. The effects of VZ after 45-dayexposure,
along with a complementary assessment of offspring embryonic
development (F1 embryos) and hatching success were monitored.
The F1 endpoints were evaluated for both non-exposed and exposed
embryos in order to distinguish effects due to parental exposure. A combination of genetic and epigenetic endpoints to study the effects
associated with mechanisms of inheritance was simultaneously
assessed. The Genotoxicity of adults and F1 embryos was assessed
by micronucleus test. The post-exposure profile alterations of global
DNA methylation in adults was the epigenetic mechanism studied.
The results were compared with our previous studies of VZ on
Physaacuta embryos whose parents were not exposed [15]. Finally
since long-term exposure was needed, actual VZ concentrations were
measured.

Material and Methods

Acute and reproduction tests

Physaacuta individuals came from our own laboratory breeding
stocks. Culture conditions are described by Sánchez-Argüello et al.
[36].

In order to stablish the exposure range of the PLC test, a prescreening
test with adults was performed at two concentrations (0.5
and 5mg/L) for 96h. Mortality and reproduction (only the number of
egg masses) were monitored in triplicate.

For the definitive test, the adults from a single cohort cultured in
our laboratory, and not previously exposed, were used for the PLC
test. Groups of ten snails (size of test animals 7±2mm) were exposed
to VZ in 60 ml of test medium using acetone (0.1%) as a solvent
carrier at 20°C, with a photoperiod of 16h:8h light/darkness. Vessels
were covered to prevent snail escaping. Maximal VZ concentration
detected in surface water was 0.0005mg/L [35]. The range of exposure
covered this measured environmental concentration and 1000 times
higher. The VZ test treatments were: Control (reconstituted test water
as used in 15); Acetone; 0.0005; 0.005; 0.05; 0.5 and 5mg/L. VZ was
added by medium renewal and after spiking nominal concentrations
on all the other days to maintain nominal concentrations. Test
medium renewal, food supply (1.2 mg/snail of Shrimps food flakes
-Sera®) and monitoring the endpoints (mortality and fecundity) were
done twice-weekly over a 45-day period. Fecundity was assessed as the
egg masses deposited per adult. The number of eggs inside each egg
mass was also counted with a stereomicroscope (Olympus SZX12).
Four replicates per treatment were used. Snails were recovered at the
end of tests for the micronuclei induction study and DNA isolation
and methylation analyses. Measures of chemical body burden were
also taken.

F1 embryo toxicity tests

Oviposition was observed on the first day of monitoring (3 days
after the test began) in all the treatments. Once a week spawned
egg masses from Oviposition were sampled and washed through a
stainless steel strainer tore move jelly-envelopment. This treatment
allows single eggs to be obtained, which are easily observed under
a stereomicroscope during the embryonic development instead of
the three-dimensional arrangement of eggs in the egg mass [15].
Single eggs were used for testing embryo toxicity. F1 eggs from each
treatment and control were divided into two batches: one batch
was transferred to the 12-well plate with reconstituted water (nonexposed
embryos), while the other batch was exposed to the same
concentrations as their parents (exposed embryos). These differences
in the exposure conditions for offspring help differentiate the effects
due only to parental exposure from the effects observed by continuous exposure (F0-adults and F1-embryos), respectively. Embryo toxicity
tests were performed in duplicate per treatment, and for both the
embryos non-exposed and exposed to VZ, in accordance with the
conditions described in Sánchez-Arguello et al. [15]. Briefly, a density
of 30 eggs/well in 4ml/well was used. The embryo toxicity test lasted
21, for this time mal formed and dead embryos were monitored daily
by stereomicroscope scoring. (Figure 1) shows differences between
normal embryos, malformed embryos and dead embryos at different
embryonic stages. Test medium was renewed twice weekly. Hatching
success was measured until the test ended.

The second, third, fourth and fifth F1 broods were also recovered
and embryo toxicity tests were conducted as described above.

Micronucleus test

Two replicates and five snails from each replicate were used for
study Genotoxicity. Shells of were gently broken between two glass
slides, and the mass was separated and immersed in 2ml HBSSBioWhittaker
® buffer (Lonza). Mechanical cell dissociation was
carried out in a glass/glass tissue grinder (Sartorius).Pieces of tissue
were separated by centrifugation (5min at 250g). The supernatant was
incubated 2h with 6 mg/ml collagenase type IA (Sigma) at 37°C, and
the cell suspension was passed through 70-μm gauze. At this point,
the same procedure described by Sanchez-Arguello et al. [15] for
micronucleus scoring with Physaacuta embryos was followed exactly,
except that the dechorionation step was excluded.

Part of the recovered egg masses of the first Oviposition was
used for testing micronuclei induction in the F1 embryos. Exposure
continued in the 12-well plates (2wells/treatment) for 1 week
under same treatment conditions as their parents. Then eggs were
treated exactly as described by Sánchez-Argüello et al. [15] for the
micronucleus test.

Preparations were coded to determine micronucleus frequencies.
For each treatment two replicates were scored and five hundred cells
were counted for each replicate, being scored a total of 1,000 cells/
treatment.

Global DNA methylation pattern analysis

DNA was extracted with the Master Pure TMDNA Purification kit (Epicentre® Biotechnologies). Restriction digestions of 1μg DNA
were performed with 10 units of either Hpa II and Msp I in buffer SL
(Sigma), and were incubated for 18h at 37°C. Hpa II and Msp I are
isoschizomers that recognize the tetranucleotide sequence 5’-CCGG-
3’but show differential cleavage sensitivity to cytosine methylation.
Digestion products were stored at -20°C until amplified by PCR.

Methylation Sensitive Amplified Polymorphism (MSAP) was
performed using two primers (MGC0: 5’-AAC CCT CAC CCT AAC
CGC GC + MGF2: 5’-AAC CCT CAC CCT AAC CCG CG) or three
primers (MGC0 + MGF2+ MLG2: 5’-AAC CCT CAC CCT AAC
CCCGG) (Stab Vida) which arbitrarily bound within the GC-rich
regions of DNA. PCR reactions were performed in a total volume of
50μl that contained: 0.5μl of each primer (0.5μM), 25μl BioMixTM
Red (Bioline) and 50mg of digested DNA. The reactions were carried
out in a thermocycler (UnoCycler, VWR). The cycling conditions
include five cycles of 30s at 94°C, 1min at 40°C, and 1.5min at 72°C
[based on 37,24].

A high-resolution display of the IRDye-labelled PCR products
was obtained using the primers labeled with 5’IRDye®-800 (Fisher
Scientific). The IRDye-labelled DNA fragments were separated into
denaturing polyacrylamide gels (16% Long Ranger®, 50% Gel Solution
(Lonza), 7M urea and 1xTBE) at 1,500 V for 2h and 30min in a Li-Cor
4300 DNA Analysis System. Fragments of 100-400 bp were visually
scored. The appearance or disappearance of amplified fragments was
considered. Fragments were scored as present (1) or absent (0). The
MSAP data were interpreted according to the review of Fulnecek and
Kovarik [38]. (Table 1) summarizes the Msp I/Hpa II profile and the
forms of the CCGG sites, which were considered [38]. The Msp I/
Hpa II profile observed in the controls was compared with the Msp I/
Hpa II profile of the treated samples, and the following interpretation
was used:

Table 1: Effect on cleavage and the expected Methylated Sensitive Amplified
Polymorphism (MSAP) profile according to Fulnecek and Kovarik [38] after
digestion with isoschizomers Msp I and Hpa II. (1) A sequence is cut and a band
is present in the electrophoretic profile of MSAP, (0) a sequence is not cut and
a band is absent.

Site

5’CCGG

GGCC5’

5’CmCGG

GGmCC5’

5’CmCGG

GGCmC5’

5’mCCGG

GGCmC5’

Msp I

1

1

0

0

Hpa II

1

0

0

0

Table 1: Effect on cleavage and the expected Methylated Sensitive Amplified
Polymorphism (MSAP) profile according to Fulnecek and Kovarik [38] after
digestion with isoschizomers Msp I and Hpa II. (1) A sequence is cut and a band
is present in the electrophoretic profile of MSAP, (0) a sequence is not cut and
a band is absent.

Chemical analysis

VZ water concentrations were determined by the HPLC analysis
with UV detection after SPE [based on 32]. Quantification was
performed using external standards (fortified water samples 500-
15ng/ml). The limits of quantification and detection were established
as15ng/ml and 5ng/ml respectively.

The adult snails collected at the end of the test were homogenized
in acetone (Ultra-Turrax). The body burden measures of VZ were
taken. The obtained extracts were analysed by GC-EC (HP-5 30m x 0.25 mm Agilent 19091J-413 column, 1.5 ml/min He and μECD). VZ
recovery was higher than 95% when fortified blank snails were used.

Statistical analysis

The fecundity, fertility and micronuclei data were compared.

Student’s t-tests were performed to analyze differences in
the fecundity (egg masses/adult) data of the pre-screening test.
Nevertheless, fecundity expressed as egg masses and cumulated
number of eggs produced per snail did not satisfy the normal
distribution criteria. In this case, a Kruskal-Wallis test, followed by a
Mann-Whitney test, was applied to study with the controls.

Hatching success data of the first Oviposition was tested for
normality. Both the F1 non-exposed and exposed embryo toxicity
data of the controls and solvents were normally distributed.
Student’s t-tests were performed to study if the solvents differed
significantly from the controls. Acetone did not produce significant
embryo toxicity. Thus the data for the solvents and treatments were
transformed to hatching percentages compared with their plate
control wells. These transformed data were normally distributed and
a two-factor ANOVA was performed. Differences due to the factor
of treatments (Acetone; 0.0005; 0.005; 0.05; 0.5 and 5mg/L) and
the factor of the exposure conditions (F1 non-exposed or exposed
embryos) were studied. Finally, a one-way ANOVA, followed by
a multiple comparison test (LSD), was applied to discern which
treatments differed from the solvent controls.

The data of the subsequent Oviposition (percentage of embryo
development completed, malformations and lethal effects) were set as
the dependent variables, and the multivariate and multifactor GLM
(General Linear Model) was used for statistical discriminations.

The micronuclei frequencies of adult snails and F1 embryos for
controls and acetone were grouped. Data fitted normal distributions
and Student’s t-tests were performed to study if the solvents differed
significantly from the controls. Acetone produced no significant
MN induction compared with the controls. Then MN frequencies
were transformed and were represented only for the solvent and
treatments in accordance with Sánchez-Argüello et al. [15]. MN data
were transformed by dividing absolute frequencies by the average of
their respective control assay, represented as MN induction above the
control levels (relative MN induction). Transformed MN data were
distributed normally and Student’s t-tests were applied to study the
differences compared with their respective solvent controls.

Statistical analyses were performed by standard procedures with
the SPSS 13.0package.

Results and Discussion

Acute and reproduction tests

Acute and reproduction tests
No mortality was observed in either the pre-screening test or
the lower treatment concentrations of the PLC test. Approximately
10% of mortality was found in the PLC test (Table 2), which is
acceptable for assessing reproduction, and is as recommended for
other standardized tests [i.e. 20% for the OECD Daphnia magna
reproduction test, [4]. The highest treatment concentration produced
a lower feeding rate and an avoiding behaviour in snails at the
beginning of the test, which disappeared during the test. This derived
ultimately in lethal effects. However, fecundity was also tested at the
highest treatment concentration (5mg/L) in two replicates, where
mortality was kept lower than 10% by replacing the dead snails from
the other two replicates. A strong and statistically significant decrease
of fecundity was observed for this highest treatment concentration
(Kruskal-Wallis test p<0.01, df=6, n=312; Mann-Whitney test
for the controls versus 5mgVZ/L p<0.01).None of the other VZ
concentrations had any effects on fecundity, although a slight rise in
Oviposition was noted, which increased according to concentration,
but was not statistically significant.

Previous research into the effects of VZ has indicated that this
fungicide may alter the reproduction of molluscs. Tillmann et al.
[34] described an earlier sexual repose of Marisa cornuarietis males
at lower VZ concentrations (0.03 and 0.1μg/L) than we used herein.
Lagadic et al. [39] pointed out the adverse effect of VZ as a result
of endocrine disruption, whose nature and amplitude probably
depended on the exposed life-stages and exposure conditions. In
their preliminary investigations into VZ, these authors observed that
the production of egg masses of L. staganalis was stimulated, but the
hatching rate lowered after exposure to 0.25mg/L. Ducrot et al. [35]
exposed young adults and adults of the pond snail Lymnaeastagnalis
to a range of VZ concentrations (0.000025; 0.00025; 0.0025; 0.025; 0.25
and 2.5mg/L) to test fecundity. While fecundity significantly reduced
in young snails at all tested concentrations, adult snails exhibited
lower fecundity and mortality during the test, but only at 2.5mg/L.
These authors concluded that age and maturation status influenced
the sensitivity of the PLC experiments. The eggs produced by young
snails were generally not fertilized, which the authors associated with
impaired the male function. Using a similar range of concentrations
for the adults of Physaacuta, we also observed effects on fecundity in
adult snails at 5mg/L.

F1 embryo toxicity tests: Fertility

The fertility of the first Oviposition (Figure 2) was statistically
affected by the factor of treatments (two-factor ANOVA; F= 3.40; p<0.05; df=5) but not by the factor of the non-exposed/exposed
conditions (Two-factor ANOVA; F= 0.32; p>0.05; df=1). The highest
treatment concentration produced statistically significant effects
compared with the controls (One-way ANOVA/LSD; F= 3.8; p<0.05;
DF=5). A sharp drop in hatching (32%) was observed when the F1
eggs were exposed during embryonic development (Figure 2), while
the hatching percentage of the F1 non-exposed eggs was higher
(54%). These results showed recovery trends for the embryos with
parental exposure when exposure was suspended during embryonic
development (Figure 2). Nevertheless, lower concentrations produced
few differences in the effects between the F1 exposed and non-exposed
embryos, as shown in (Figure 3) for the 0.5 mg/L treatment. This effect
was observed in the following Oviposition (second, third, fourth and fifth F1 broods) for all the treatments below 5 mg/L. A similar slope of
the curves for effects on embryos (normal embryonary development,
mortality and malformations) in both F1 exposed and non-exposed
embryos is observed in (Figure 4). Although the variability of data for
F1 embryo toxicity was higher in the exposed embryos than in the
non-exposed ones, the subsequent exposure conditions did not cause
statistical differences for any variable (GLM; F= 0.66; p>0.05). The
larger the brood number, the less successful hatching. The F1 hatching
percentage lowered in the controls from 30% to 0% between second
and fifth brood, respectively. Normal embryonary development
was affected by both brood number (one-way ANOVA; F=23.04;
p<0.05; DF=2) and treatments (one-way ANOVA; F= 4.65; p<0.05;
DF=6). Malformations increased depending on brood number (oneway
ANOVA; F=25.95; p<0.05; DF=2) and treatments (one-way
ANOVA; F=3.12; p<0.05; DF=6). Nevertheless, lethal effects were
affected by treatments (one-way ANOVA; F=3.01; p<0.05; DF=6)
regardless of brood number (one-way ANOVA; F= 2.19; p<0.05;
DF= 2). Therefore, lethal effects were caused mainly by treatments,
while malformations were attributed to the quality of broods. These
findings lead to the conclusion that for the next PLC experiments,
the first brood would suffice to study fertility since its viability was
the best.

Figure 2: Percentage of hatching for offspring (F1 embryos) compared with
the control. The hatching percentages for solvents came close to 70% (67.6%
± 17.4 for exposed and 71.3% ± 20.6 for non-exposed). It was not possible
to represent them because the logarithmic scale was used (treatments=0).
Statistically analysis after a one-way ANOVA, followed by a multiple
comparison test (LSD), showed which treatments differed from the solvents
(*p<0.05).

Figure 2: Percentage of hatching for offspring (F1 embryos) compared with
the control. The hatching percentages for solvents came close to 70% (67.6%
± 17.4 for exposed and 71.3% ± 20.6 for non-exposed). It was not possible
to represent them because the logarithmic scale was used (treatments=0).
Statistically analysis after a one-way ANOVA, followed by a multiple
comparison test (LSD), showed which treatments differed from the solvents
(*p<0.05).

Figure 3: Embryonic effects (hatching, embryo development completed,
lethal effects and malformations) for both the F1 non-exposed and exposed
embryos after exposure to 5 and 0.5mg/L.

Figure 3: Embryonic effects (hatching, embryo development completed,
lethal effects and malformations) for both the F1 non-exposed and exposed
embryos after exposure to 5 and 0.5mg/L.

Figure 4: The embryonic effects (embryo development completed, lethal effects and malformations) observed in the second, third, fourth and fifth Oviposition after
the PLC tests. Both curves of the F1 non-exposed and exposed embryos showed similar slopes but higher uniformity was observed for the non-exposed embryos.

Figure 4: The embryonic effects (embryo development completed, lethal effects and malformations) observed in the second, third, fourth and fifth Oviposition after
the PLC tests. Both curves of the F1 non-exposed and exposed embryos showed similar slopes but higher uniformity was observed for the non-exposed embryos.

Lethal effects and decreased fecundity, both observed after
adult VZ exposure to 5mg/L, also correlated with the lower fertility
(hatching success) of both exposed and non-exposed F1 embryos.
The embryo toxicity of VZ for this study did not correlate with our
previous research since this treatment did not produce embryo toxicity
for those eggs without parental exposure. More than 70% of the
hatchability for the embryos without parental exposure was observed
at 5mgVZ/L [15]. These differences in response sensitivity indicated
an increasing sensitivity of embryos in this study (both exposed and
non-exposed during embryonic development), probably due to the
chronically stressed parents. Other studies conducted with freshwater
snails have reported latent effects being expressed one generation later
[40,41]. Oliveira-Filho et al. [42] evaluated the effects on the fecundity
of mature F0 and F1 freshwater snails (Biomphalariatenagophila)
and developmental toxicity in F1 and F2 embryos after exposure
to endosulfan and ethanol. Their results indicated that while the
reproductive effects of endosulfan did not apparently change as
exposure extended to successive generations, ethanol effects became
even more marked in the subsequent generation. The effect observed
at 5mg/L herein, which did not correlate with the previous study [15]
showed that early life stages (embryos) cannot be as sensitive as the
adults of Physaacuta after chronic exposure. These results disagree
that the short-term test with early life stages often shows the same
effect concentrations as those observed with less sensitive life stages
(i.e. adults) after long-term exposure [43]. Seeland et al [18] also
observed less sensitivity of P acuta embryos than juveniles, which can
be explained by egg integument acting against harmful environmental
influences. Munley et al. [17] found that growth effects were
predictive of reproduction effects when comparing a 28-day early
life-stage test and a Full Life-Cycle (FLC) test on Lymnaeastagnalis,
although the embryonic growth of F1 organisms was not included in
the comparative data analysis.

Micronucleus tests

The adults exposed to 0.5 and 5mg/L showed statistically significant micronuclei induction. Although fungicide VZ has been
ranked as essentially non-genotoxic [44,45] very few reports have
indicated that it induces Genotoxicity (i.e. micronuclei, chromosome
aberration and sister chromatid exchange) in mammalian cells in
vivo and in vitro [46,47,31]. The highest concentration (5 mg/L)
also produced a statistically significant induction of micronuclei
in embryos without parental exposure, as previously observed by
Sánchez-Argüello et al [15]. All the other treatments did not produce
statistically significant MN induction in F1 embryos herein (Figure
5). Further research to make comparisons is required since the 0.5
mg/L treatment was not tested in embryos without parental exposure
[15] and also because it was not possible to test 5mg/L in F1 embryos
as it were embryo toxic.

Figure 5: Relative induction of micronuclei in the cells isolated from adult
snails and the F1 exposed embryos. Data are shown as mean ± std (n=2;
1000 cells/n). Statistically significant differences after the Student’s t-tests
referred to the comparison made between treatments and solvents (*p<0.05;
**p<0.01).

Figure 5: Relative induction of micronuclei in the cells isolated from adult
snails and the F1 exposed embryos. Data are shown as mean ± std (n=2;
1000 cells/n). Statistically significant differences after the Student’s t-tests
referred to the comparison made between treatments and solvents (*p<0.05;
**p<0.01).

Effects on DNA methylation

MSAP-PCR proved valuable for detecting that Physaacuta
displays DNA methylation for the first time because we observed
differences between profiles Msp I and Hpa II. Some modifications
to the global DNA methylation pattern were also observed among
the different treatments. Both methylation and demethylation events
took place in treatments (Figure 6) compared with the control.
Studies into the epigenetic effects of VZ exposure have focused on
mammals. VZ induces both methylation and demethylation events
in 25 regions in the rat genome [30]. A wide promoter analysis has identified 66 mouse promoters with 68 differential DNA methylation
regions [48]. Studies on epigenetic biomarkers with exposed
invertebrates are rare. When it comes to DNA methylation, insects
have been the most widely studied invertebrate group [23]. One of
the earliest investigations that studied the role of DNA methylation in
invertebrate gene expression reported differences in patterns between
insecticide-resistant aphid clones and those that had lost resistance
[49]. Del Gaudio et al. [50] showed for the first time that the DNA of a
polychaete annelid marine worm is methylated and that methylation is lower in adults than in embryos or sperm DNA. Krauss et al. [51]
demonstrated that DNA methylation can play a key role in gene
activity regulation during development in walking stick insects.
Vandegehuchte et al. [33] demonstrated a reduction in the overall
DNA methylation of Daphnia magna after exposure to 430μg/L of
VZ, and also revealed that toxicant exposure can not only perturb
the DNA methylation status of water fleas, but be transferred to
two subsequent non-exposed generations. Very little work has been
done to investigate DNA methylation in molluscs. The first evidence
that supported a regulatory role of intragenic DNA methylation in
invertebrates was done in a pacific oyster. The authors of this work
suggested that DNA methylation in molluscs perform regulatory
functions, including those involved in stress and environmental
responses [52]. Their research showed an alteration to the DNA
methylation profiles of treatments compared with the control, as
expressed qualitatively in the events of both demethylation (five sites)
and methylation (two sites).

Figure 6: The methylation-sensitive arbitrarily primed PCR (MSAP-PCR)
analysis of the DNA digested with MspI and HpaII in polyacrylamide gels. Two
methylation profiles were obtained using two primers (a) and three primers
(b). Circles show demethylation events and squares denote methylation
events, according to Fulnecek and Kovarik (2014).

Figure 6: The methylation-sensitive arbitrarily primed PCR (MSAP-PCR)
analysis of the DNA digested with MspI and HpaII in polyacrylamide gels. Two
methylation profiles were obtained using two primers (a) and three primers
(b). Circles show demethylation events and squares denote methylation
events, according to Fulnecek and Kovarik (2014).

Chemical analysis

The exposure regime allowed concentrations to remain close to the
nominal one during the experiment, as shown by the measurements
taken before and immediately after medium renewal (Table3a).
Previous experiments showed that VZ disappears rapidly in water
(half-life of 10.8 h, 15] and is probably degraded to metabolites, but
the presence of food could have conditioned its behaviour in this
experiment. VZ bioavailability was expected to increase through
dietary exposure since its log Kow is 3 [53]. VZ was measured in
the snails recovered at the end of the PLC test (Table 3b). VZ might
be adsorbed to living organisms but may also be bioconcentrated/
bioaccumulated by snails. VZ body burden peaked at 0.5 mg/L since
similar concentrations in snails were observed for highest treatment
concentration.

Table 3: Actual VZ concentrations:
a) Water concentrations before and after medium renewal.
b) VZ body burden at the end of the reproduction test. No measured (nm).

a) VZ water concentrations (mg/L)

b) VZ body burden (μg/g)

Nominal

Before medium renewal

After medium renewal

0.0005

0.0002

0.0002

0.0009

0.0009

0.049

0.076

0.005

0.0050

0.0052

0.0032

0.0039

0.206

0.600

0.05

0.0578

0.0625

0.0558

0.0691

4.72

4.94

0.5

0.4072

0.4225

0.5558

0.5166

18.60

25.46

5

3.909

3.370

5.336

nm

22.95

nm

Table 3: Actual VZ concentrations:
a) Water concentrations before and after medium renewal.
b) VZ body burden at the end of the reproduction test. No measured (nm).

Conclusion

This study showed that VZ affected the fecundity of Physaacuta, but
only when effects on survival after long-term exposure were observed
(5mg/L). The lower concentrations neither produced mortality nor
affected fecundity. Similarly, fertility was affected only at 5mg/L. This
concentration was not embryotoxic in a previous study conducted
with embryos without parental exposure [15], which indicates that
embryos were more sensitive when parents were exposed. Therefore,
it can be concluded that a reproduction assessment in freshwater
invertebrates should include both fecundity (eggs production) and
fertility (offspring viability). Additionally comparisons between exposed and non-exposed offspring help indicate whether damage
due to parental exposure can be maintained or recovered.

The highest treatment concentration (5mg/L) also resulted in
Genotoxicity for adults. Recent studies have offered lines of evidence
that MN formation is induced epigenetically. Luzhna et al. [25]
have indicated that global DNA methylation loss correlates with
MN formation. Demethylation was observed4 times at 5mg/L. This
study provides information on the methylation pattern in a species
representing a phylum where almost no information regarding this
topic is available at the moment and show that a known epigenetic
toxicant can cause alterations. Although data on the epigenetics
of molluscs in an ecotoxicological context are scarce, this study
shows that they are suitable for continuing with this research line.
Both genetic and epigenetic molecular mechanisms are associated.
The combination of genetic/epigenetic biomarkers for establishing
mechanisms of inheritable molecular basis is recommended.

Reproduction success is ultimately the relevant endpoint
to address effects at the population level. The linkage between
reproduction effects and mechanisms capable of inducing heritable
changes could broaden our understanding of pollutant impacts.